Acute waterborne cadmium toxicity in the estuarine pulmonate mud snail, Amphibola crenata

Ecotoxicology and Environmental Safety 158 (2018) 274–283

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Acute waterborne cadmium toxicity in the estuarine pulmonate mud snail, Amphibola crenata

T

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Nuwan A.L. De Silvaa, , Islay D. Marsdena, Sally Gawb, Chris N. Gloverc,d a

School of Biological Sciences, University of Canterbury, New Zealand School of Physical and Chemical Sciences, University of Canterbury, New Zealand c Faculty of Science and Technology and Athabasca River Basin Research Institute, Athabasca, University, Athabasca, Alberta, Canada d Department of Biological Sciences, University of Alberta, Edmonton, Canada b

A R T I C LE I N FO

A B S T R A C T

Keywords: Bioenergetics Cadmium Gastropod Trace metal Oxidative stress Protein metabolism

Freshwater pulmonate snails are sensitive to trace metals, but to date, the sensitivity of estuarine pulmonate snails to these important environmental toxicants is undescribed. Using the estuarine mud snail Amphibola crenata, effects of a 48-h exposure to waterborne cadmium (Cd) were investigated. The 48-h median lethal concentration (LC50) was 50.4 mg L−1, a value higher than that previously reported for any gastropod mollusc. Cadmium levels in the tissues of mud snails were highest in the viscera (digestive gland and gonad), with the foot muscle and remaining tissue compartment (kidney, mantle, remaining digestive tissues and heart) displaying significantly lower concentrations. Over a Cd exposure concentration range of 0–32 mg L−1, Amphibola exhibited reduced oxygen consumption and elevated ammonia excretion in response to increasing Cd, the latter effect likely reflecting a switch to protein metabolism. This finding was supported by a declining oxygen: nitrogen ratio (O:N) as exposure Cd concentration increased. Other energy imbalances were noted, with a decrease in tissue glycogen (an effect strongly correlated with Cd burden in the viscera and foot muscle) and an elevated haemolymph glucose observed. An increase in catalase activity in the visceral tissues was recorded, suggestive of an effect of Cd on oxidative stress. The magnitude of this effect was correlated with tissue Cd burden. The induction of antioxidant defence mechanisms likely prevented an increase in levels of lipid peroxidation, which were unchanged relative to Cd exposure concentration in all measured tissues.

1. Introduction Estuaries are demanding environments, exposing the biota therein to extreme fluctuations in salinity, dissolved oxygen, temperature, and nutrients (e.g. Hubertz and Cahoon, 1999), all factors that threaten organism homeostasis. Estuaries are also sinks for contaminants, which further challenge health and survival. This is especially true of estuaries located near urban centres, which may receive domestic, industrial and agricultural effluents, either through direct inputs, or via the river systems that drain into them (Matthiessen and Law, 2002). Furthermore, a variety of physical and chemical factors can effectively trap contaminants within estuaries (Ridgway and Shimmield, 2002), exposing estuarine biota to potentially harmful concentrations of a diverse range of toxicants. Among the contaminants of particular concern in estuarine settings are trace metals. These are environmentally persistent and can cause a variety of toxic effects in the exposed animal (e.g. Chandurvelan et al., 2015). Sub-lethal (e.g. biochemical and physiological) effects can lead

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to changes in organism health and fitness, eventually resulting in changes at the population, community and ecosystem level (e.g. Lagadic et al., 1994; Marsden and Swinscoe, 2014). Among trace metals, cadmium (Cd) is of significant interest. This highly toxic, non-essential trace element, is distributed ubiquitously in aquatic environments. It is released into the environment by both anthropogenic (e.g. municipal wastewater, fossil fuel combustion, metal processing, agricultural runoff; Eisler, 1985; Butler and Timperley, 1996) and natural sources, such as weathering of Cd-rich geology and volcanism (Eisler, 1985). Although the average concentration of Cd in seawater is around 0.1 μg L−1 (Korte, 1983), marine sediment concentrations can be as high as 80 μg g−1 (Böning et al., 2004). Estuarine Cd contamination has received significant recent attention in New Zealand (e.g. Chandurvelan et al., 2015, 2016), owing in part to concerns regarding Cd contamination of superphosphate fertilisers applied extensively to nearcoastal lowland streams (Butler and Timperley, 1996; McDowell, 2010), and run-off from mine tailings that can result in river Cd concentrations as high as 800 μg L−1 (Craw et al., 2005). Cadmium can

Correspondence to: School of Biological Sciences, University of Canterbury, Private Bag 4800, Christchurch 8140, New Zealand. E-mail addresses: [email protected] (N.A.L. De Silva), [email protected] (I.D. Marsden), [email protected] (S. Gaw), [email protected] (C.N. Glover).

https://doi.org/10.1016/j.ecoenv.2018.04.041 Received 31 January 2018; Received in revised form 16 April 2018; Accepted 19 April 2018 0147-6513/ © 2018 Elsevier Inc. All rights reserved.

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ratio (O:N), which have previously been shown to be impacted by Cd in aquatic biota (e.g. Barbieri and Paes, 2011; Chandurvelan et al., 2012, 2017). In addition, biochemical correlates of energy use (tissue glycogen, haemolymph glucose) were measured. Biochemical analysis also included measures of oxidative stress, based on the ability of Cd to induce reactive oxygen species (ROS), likely via displacement of redox active metals from cofactor binding sites (Nair et al., 2013). Cadmium has also been shown to impair the function of enzymes that scavenge ROS, and thus generate oxidative damage indirectly (e.g. Chandran et al., 2005; Chandurvelan et al., 2013; McRae et al., 2018). Overall, the objectives of the present study were to determine toxicological, physiological and biochemical responses of A. crenata to acute Cd exposure, and to investigate Cd bioaccumulation in different tissues as a function of exposure levels.

cause a wide variety of toxicological impacts on aquatic organisms including changes in biochemical pathways associated with oxidative stress, altered energy metabolism, impaired calcium homeostasis, perturbed growth and reproduction, and at sufficiently high exposure concentrations, death (e.g. Eisler, 1985; McGeer et al., 2012). One group of aquatic organisms that have been reported as displaying high sensitivity to trace metals (Grosell et al., 2006; Brix et al., 2011), including Cd (Das and Khangarot, 2010), are the freshwater pulmonate gastropods. For example, Lymnaea stagnalis is highly susceptible to copper toxicity, stemming from impairments in ion regulation that affect shell development (Brix et al., 2011), while inhibition of feeding and growth are observed at Cd concentrations as low as 32 μg L−1 in L. luteola (Das and Khangarot, 2010). It is, however, important to note that not all studies support the conclusion of high sensitivity of freshwater pulmonate snails to trace metals. A study on copper toxicity comparing pulmonate snails with hydrobiid snails, found that toxicity was similar in both groups (Besser et al., 2016). The difference in the sensitivity of pulmonate snails in this study relative to other work, was attributed to the authors use of a distinct age class of animals (Besser et al., 2016). However, this finding of relatively high tolerance is consistent with Cd toxicity studies in terrestrial pulmonate snails (Chabicovsky et al., 2004). It is intriguing that, irrespective of sensitivity, both freshwater and terrestrial pulmonates have an excellent capacity for Cd bioaccumulation (Chabicovsky et al., 2004; Das and Khangarot, 2010). For example, bioconcentration factors of more than 6000 have been reported in L. palustris exposed to 160 μg L−1 waterborne Cd for four weeks (Das and Khangarot, 2010). However, we are unaware of any published reports that detail the sensitivity of marine or estuarine pulmonate snails to Cd, or that delineate their capacity for Cd bioaccumulation. In the current study, the mortality, bioaccumulation, biochemical and physiological impacts of waterborne Cd exposure were examined in the estuarine pulmonate mud snail, Amphibola crenata. This species is widely distributed throughout New Zealand and is found in both relatively clean and contaminated areas (Marsden and Baharuddin, 2015). It has a sedentary lifestyle, is abundant year-round, easy to collect, and is highly tolerant to extremes of temperature, aerial exposure and desiccation (Shumway, 1981; Shumway and Marsden, 1982). Little is known regarding the sensitivity of this species to metal contaminants, however it has been shown that those snails inhabiting settings with high sediment metals display a decreased condition index (Marsden and Baharuddin, 2015). Our studies were conducted in the laboratory under acute exposure conditions. Laboratory-based testing overcomes some of the practical difficulties in determining the effects of individual stressors in the field, and thus is an essential tool for understanding the mechanisms of contaminant impacts in organisms. An understanding of toxic mechanisms is itself essential for development of predictive models, allowing extrapolation between different species on the basis of shared pathways of impact. Mechanistic data are also increasingly important for the development of regulatory tools (e.g. Biotic Ligand Model), which facilitate site-specific determination of toxicity based on knowledge of water chemistry, and mechanisms of uptake and toxicity (Di Toro et al., 2001). Studies performed over acute time-frames provide rapid and reproducible estimates of the effects of contaminants on biota, and are critical to the development of acute-to-chronic toxicity ratios, which are the basis of many water quality criteria (e.g. Shuhaimi-Othman et al., 2013). Moreover, when coupled with measurement of mechanistic endpoints, these tests provide an overview of toxic responses in an organism (Chandurvelan et al., 2012), and may help in the future selection of biomarkers and bioindicator species for environmental monitoring (e.g. Shuhaimi-Othman et al., 2012). In the current study, biochemical and physiological endpoints were used to characterise the mechanisms by which Cd exerts its effects. These included physiological indices of metabolic impairment such as oxygen consumption, ammonia excretion and the oxygen to nitrogen

2. Materials and methods 2.1. Sample collection and maintenance Adult mud snails (> 18 mm) were collected during December 2015 from the mouth of the Avon-Heathcote Estuary/Ihutai (S 43°33.136', E 172°44.709') in Canterbury, New Zealand. Recently measured sediment Cd concentrations near the site of collection were 0.1 μg g dry weight−1 (Chandurvelan et al., 2016). Snails were then transported to the aquarium at the University of Canterbury, where they were washed thoroughly with filtered natural seawater, and attached algae and mud was scraped from the shells. Snails were then transferred into a holding tank containing 2 L of 20 ppt filtered natural seawater (pH 7.6; Na 325 mmol L−1, Ca 9 mmol L−1; Mg 37 mmol L−1; K 8 mmol L−1; Cl 360 mmol L−1; dissolved organic carbon 0.3 mg C L−1) made by diluting Lyttelton Harbour seawater with City of Christchurch artesian well water, and acclimated for 48 h prior to the experiments, under constant temperature (15 ± 0.5 °C) and photoperiod (12 h light: 12 h dark). Snails were not fed during acclimation. 2.2. Determining median lethal concentration (LC50) Toxicity testing was carried out in constantly-aerated acid-washed 1.5 L polypropylene containers at 15 °C and 20 ppt salinity (natural SW as described above) at six Cd exposure concentrations (nominally: 0, 8, 16, 32, 64, and 128 mg Cd L−1), achieved by the dilution of a stock solution of 10 g L−1 Cd (as CdCl2.2½H2O). Each exposure concentration was replicated 6 times, with 15 individual mud snails per replicate. Snails were assigned randomly to the different treatments, and were not fed during the 48-h exposure period. A 15-mL water sample was taken from each treatment at time 0 and 48 h, by filtering through a Millex 0.45 μm filter (Millipore Ltd, Cork, Ireland), with these samples subsequently acidified to pH < 2 using 70% ultrapure HNO3. Acidified samples were then diluted 30 × with milli-Q water and stored at 4 °C until analysed by Atomic Absorption Spectroscopy (see below). Mortality was assessed every 12-h throughout the experiment period, with death defined as the point when immobile mud snails failed to respond to probing using forceps. The LC50 values were calculated based on measured Cd exposure concentrations. 2.3. Bioaccumulation, biochemistry and physiology Surviving individuals from the 0, 8, 16 and 32 mg Cd L−1 exposure concentrations used to determine the LC50 were then used for assessment of tissue Cd content, and physiological and biochemical responses to Cd exposure. While these exposure concentrations are significantly elevated relative to those likely to be experienced in natural settings, they provide a means of exploring the mechanisms by which Cd exerts toxic effects on Amphibola. Immediately following the 48-h exposure, two groups of snails were subjected to physiological assays (see below; each n = 6), while a 275

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2.3.2. Biochemical endpoints Tissue catalase activity was determined by an assay utilising the decomposition rate of hydrogen peroxide (H2O2; Chandurvelan et al., 2013; McRae et al., 2016). The tissue samples were homogenised in 800 μL ice-cold homogenisation buffer (100 mM Trizma base, 2 mM EDTA, 5 mM MgCl2·6H2O; pH = 7). Of this homogenate, 200 μL was used for lipid peroxidation assessment (see below). The remaining homogenate (600 μL) was centrifugation at 30,000 × g for 10 min at 4 °C. The resulting supernatant was diluted 30 × with the homogenisation buffer, and the reaction was initiated by the addition of H2O2, with the decline in absorbance at 240 nm measured in a microplate reader at 25 °C (UV star; Greiner Bio-One). Supernatant protein concentration was measured via a Bradford assay (Bradford, 1976), using a standard curve based on bovine serum albumin. Catalase activity was expressed as μmol mg protein−1 min−1. Tissue lipid peroxidation was measured using a commercial lipid peroxidation assay kit (Sigma Aldrich, MAK085), according to manufacturer instructions. This method is based on the reaction of malondialdehyde (MDA) with thiobarbituric acid (TBA) to form a coloured product proportional to the MDA present in the sample. The assay was initiated by adding 300 μL of MDA lysis buffer and 3 μL of butylated hydroxytoluene (BHT) to the homogenate (200 μL; see above). This mixture was centrifuged at 13,000 × g for 10 min. To 200 μL of the resulting supernatant, 600 μL of TBA solution was added, with this solution incubated at 95 °C for 1 h. Once samples cooled, 200 μL was transferred to a 96-well plate reader and absorbance was measured at 532 nm. The amount of tissue lipid peroxidation was expressed as μmol MDA mg protein−1, where protein levels were derived from the Bradford assay (see above). Tissue glycogen concentration was determined by a coupled enzyme assay (Sigma Aldrich, MAK016). The tissue samples were homogenised in ice-cold deionised water and boiled for 5 min to deactivate glycolytic enzymes. Samples were then centrifuged at 13,000 ×g and the resulting supernatants were diluted 100 × using deionised water. After adding hydrolysis buffer and hydrolysis enzyme mix, samples were incubated at room temperature for 30 min. Master reaction mix, including development buffer, development enzyme mix and fluorescent peroxidase substrate, was added to each sample and absorbance was measured at 570 nm using a 96-well plate reader. To quantify the glycogen concentration, a calibration curve was performed using a commercially available standard glycogen (Sigma Aldrich). Glycogen level was expressed as mg g wet weight−1. The glucose concentrations in snail haemolymph were quantified using a commercial glucose assay kit (Sigma Aldrich, GAHK-20), according to manufacturer instructions. Haemolymph protein concentration was measured via the Bradford method (Bradford, 1976) and expressed as mg mL−1.

separate group of animals was sampled for tissue Cd bioaccumulation (n = 6) and for biochemical assays (n = 6). Each n value represents an individual snail from a separate exposure concentration replicate, to avoid pseudoreplication. For animals destined for biochemical assays, haemolymph samples were collected from the foot muscle sinus using a 21-gauge needle and syringe. Then for both accumulation and biochemistry endpoints, snails were blotted dry removing any adherent mucus, and dissected into three anatomical regions: gonad and digestive gland, foot muscle, and remaining tissues (i.e. kidney, mantle, remaining digestive tissues and heart). The gonad and digestive gland (hereafter denoted as viscera) were grouped as one tissue due to practical difficulties to separate them completely. All tissue samples were immediately stored at − 80 °C until further analysis. Where appropriate, all biochemical biomarker responses were expressed in terms of wet tissue weight. For snails subjected to physiological studies, all soft tissues were removed and dried to a constant weight in an oven at 60 °C for 3 days. Dissected tissues for Cd analysis were treated similarly, and subsequently all values for physiological biomarkers and tissue Cd content were expressed in terms of dry tissue weight of the mud snails. 2.3.1. Physiological endpoints Closed system aquatic respirometry was used to measure the rate of oxygen consumption of individual mud snails (n = 6), following a protocol described by Chandurvelan and colleagues (2012). Glass respirometry chambers (100-mL) were filled with oxygen-saturated 20 ppt seawater and maintained in a 15 °C water bath overnight. Preweighed mud snails were then placed in each chamber, which was sealed tightly with a rubber bung. This bung had a syringe port which facilitated the sampling of the water. A control chamber without a mud snail was also maintained, to account for any microbial contribution to oxygen consumption. Water samples were taken at time 0 and 1 h and the decline in oxygen partial pressure (PO2) inside the chamber was recorded using an oxygen electrode (Strathkelvin 1302), which was calibrated before each measurement in oxygen-saturated 20 ppt water and a sodium sulfite solution. The PO2 was recorded via a PowerLab (ADInstruments, Waverly, Australia) data recording system. The oxygen consumption rate was calculated using following equation and expressed as μmol O2 g−1 h−1:

O2 consumption rate =

∆PO2 × C × V W×t

where, ∆PO2 is the change in PO2 inside the chamber, C is O2 solubility (μmol O2 L¯1) at 15 °C and 20 ppt, V is the volume of water corrected for snail mass (L), W is the snail soft tissue dry weight (g) and t is the time (h). To determine ammonia excretion, a separate group of individual mud snails (n = 6) was transferred into 50-mL plastic containers filled with 20 mL of 20 ppt filtered seawater. Containers were covered with aluminium foil and left for 6 h at 15 °C. A blank without a mud snail was also maintained. Initial and final ammonia concentrations were measured using a salicylate-based ammonia assay (Charan-Dixon et al., 2017). Ammonia excretion (μg g−1 h−1) was then calculated using following equation:

Ammonia excretion =

2.4. Water and tissue Cd concentrations Acidified and filtered seawater samples (pH < 2) taken from the Cd exposures, were diluted 30 × with 2% HNO3. Diluted samples were then analysed by Atomic Absorption Spectroscopy (AAS, Varian; 220FS) at wavelength of 326.1 nm. To quantify the Cd concentration, a calibration curve was performed using a commercially-available Cd standard (Fluka Analytical). The limit of detection for this approach was 1 μg L¯¹ . The reported measured Cd exposure concentrations represent the mean values of initial and final water samples for each exposure replicate, which were then averaged across all replicates. Measured Cd concentration data were then subjected to Cd speciation analysis to determine the concentration of bioavailable Cd (i.e. Cd2+). This analysis was performed using the water chemistry described in Section 2.1 and the Visual MINTEQ geochemical modelling program (ver. 3.1; Gustafsson, 2012). Cadmium bioaccumulation in snail tissues was determined using inductively coupled plasma mass spectrometry (ICP-MS, Agilent-

∆C NH4+ × V 1000×t × W

where, ∆C NH4+ is the change in NH4+ concentration (μg L¯1), V is the volume of water (L), t is the time (h), and W is the dry weight of the snail soft tissue (g). The O:N ratio was then estimated from oxygen consumption and ammonia excretion values as follows (Chandurvelan et al., 2012):

O : Nratio =

(O2 mgh−1)/16 (NH4+mgh−1)/14

As oxygen consumption and ammonia excretion were measured in separate animals, only mean values were used. 276

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The tissue Cd concentration of the control mud snail group was low (< 0.8 μg g dry wt.−1) in all tissues. However, in snails exposed to Cd, increases in tissue Cd burden were observed, with a two-way ANOVA showing significant effects of Cd exposure concentration (p < 0.01), tissue (p < 0.01), and also a significant interaction between these two factors (p < 0.01; Fig. 1). In all Cd exposure concentrations, the viscera (gonad and digestive gland) displayed a significantly higher Cd accumulation than that found in the foot muscle and remaining tissues (i.e. kidney, mantle, remaining digestive tissues and heart). However, in the viscera and the foot, there was no significant increase in tissue burden when increasing Cd exposure concentration from 8 to 32 mg L−1. This is in contrast to the “remaining tissues”, where a significant increase in tissue Cd was seen at 32 mg Cd L−1. The relationship between Cd exposure concentration and tissue Cd accumulation was best described by a hyperbolic relationship for viscera (R2 = 0.645, p < 0.001) and remaining tissues (R2 = 0.811, p < 0.001), and a linear relationship for foot muscle (R2 = 0.822, p < 0.001; Table 2). Relative to the control, oxygen consumption rates for A. crenata exposed to 8, 16 and 32 mg Cd L−1 decreased significantly by 56%, 76% and 92% (p < 0.05; Fig. 2A). There were, however, no significant differences in oxygen consumption between Cd-exposed snails. Ammonia excretion increased significantly (by 68% and 84%, respectively) in animals exposed to 16 and 32 mg Cd L−1 (p < 0.05; Fig. 2B). The responses of both of these physiological endpoints were strongly and significantly correlated to Cd exposure concentrations (oxygen consumption: R2 = 0.659, p < 0.01; ammonia excretion: R2 = 0.576, p < 0.01; Table 3). Determination of oxygen consumption and ammonia excretion allowed the calculation of the O:N ratio for A. crenata. These data showed that molar ratio of oxygen consumed to ammonia excreted declined with increasing Cd exposure concentration, with values of 17, 7 and 2 reported at Cd exposure concentrations of 8, 16 and 32 mg Cd L−1, respectively. The O:N in control conditions was 50 (Fig. 2C). As oxygen consumption and ammonia excretion were determined in separate snails, these data represent group mean values, and were not able to be statistically compared. Analysis of the effect of Cd exposure on tissue catalase activity showed a significant effect of tissue type (p < 0.01), Cd exposure concentration (p < 0.01) and also a significant effect of the interaction between these two factors (p < 0.05). Post-hoc analysis showed that catalase activity in the viscera of mud snails increased with Cd exposure concentration, and was significantly higher than that in the other tissues. Foot muscle and the remaining tissues did not display a significant induction of catalase activity over the range of concentrations tested (Fig. 3A). The increase in catalase response in the viscera correlated strongly with the viscera Cd burden (R2 = 0.584, p < 0.01; Table 3). A two-way ANOVA highlighted that lipid peroxidation was not significantly impacted by tissue type (p = 0.07), Cd exposure concentration (p = 0.09) or interaction between these two factors (p = 0.27) (Fig. 3B). Acute Cd exposure significantly altered tissue glycogen levels in A. crenata. A significant effect of tissue type (p < 0.01) and Cd exposure concentration (p < 0.01) were noted, but no significant effect of the interaction between these two factors (p = 0.11) was seen. In both viscera and foot muscle, tissue glycogen significantly declined at 8 mg Cd L−1, relative to control values, although a further increase in Cd exposure concentration did not result in further significant reduction (Fig. 4A). The reduction was 36%, 47% and 62% in the viscera, while foot muscle glycogen declined by 45%, 53% and 61% for 8,16 and 32 mg Cd L−1 treatments, respectively. The glycogen levels in the remaining tissues were statistically similar to the control group for all Cd exposures. The trend in viscera (R2 = 0.338, p < 0.01) and foot muscle (R2 = 0.568, p < 0.01) glycogen concentrations correlated significantly with Cd tissue burdens in these tissues (Table 3). Relative to control animals, no significant changes in haemolymph glucose level were observed in mud snails exposed to 8 mg Cd L−1 (Fig. 4B). In contrast, significantly increased levels of haemolymph

7500cx), using a modification of the method described by Chandurvelan et al. (2012). Dried tissue samples were acid digested overnight using 0.25 mL of 70% ultrapure HNO3, before being heated at 85 °C for 1 h. The samples were diluted 5 × with 2% HNO3 and analysed by the ICP-MS for Cd. Quality assurance/quality control was achieved by analysing blanks and replicates (n = 2) of a certified mussel tissue standard reference material (SRM 2976; National Institute of Standards and Technology, US). The mean ± SD recovery of the SRM for Cd was 89 ± 8%. Tissue Cd concentration was expressed as μg g dry wt.−1. The limit of detection for tissue Cd using this method was 0.02 μg g−1. 2.5. Statistical analysis All data were processed using R statistical software (R version 3.0.2). Median lethal concentration (LC50) and 95% confidence interval values for A. crenata were calculated using probit analysis. For tissue Cd bioaccumulation, catalase activity, lipid peroxidation and glycogen levels, significant effects of Cd treatment, tissue type, and the interaction between these two factors, were determined with a two-way ANOVA followed by a Tukey's post hoc test. All other physiological and biochemical responses were analysed via one-way ANOVA followed by post hoc Tukey's test. All data were tested for, and passed, normality and homogeneity of variance assessments using Shapiro–Wilk and Fligner–Killeen tests before being analysed by parametric analyses. Regression analysis was used to determine the relationship between Cd bioaccumulation and Cd exposure concentrations, using either a linear or hyperbolic curve-fit and applied to individual, not mean, data. The relationship which fitted the data the best (i.e. highest R2 value) is reported. The goodness of fit of a linear relationship between biochemical/physiological endpoints and Cd exposure concentration (for haemolymph measures and physiological endpoints) or tissue Cd concentration (for measures in viscera, foot or remaining tissues) was also determined. All data, except O:N ratio (overall mean value only; see Section 2.3.1), are presented as mean ± SEM, unless otherwise stated. A value of p < 0.05 was considered statistically significant. 3. Results Table 1 presents the nominal and measured concentrations of Cd for each exposure treatment. These data show that measured values were close to nominal concentrations, and that there was little variation between replicates. In all concentrations, 6.6% of Cd was present in the dissolved ionic form (Cd2+). No mortality occurred in control animals throughout the experiment. In fact, mortality first occurred at a waterborne Cd concentration of 16 mg L−1, with complete mortality observed at 128 mg Cd L−1. The 48-h median lethal concentration (LC50) of A. crenata for Cd was 50.4 mg L−1 with lower and upper 95% confidence limits of 44.1 and 56.7 mg L−1, respectively. When based on the predicted Cd free ion (Cd2+) concentration, the 48-h LC50 value was 3.33 mg L−1. Table 1 Nominal Cd, measured Cd, and predicted Cd2+ (Visual MINTEQ) concentrations in exposure solutions. Nominal and measured values are expressed as mean ± SD of 6 replicates. Nominal concentration (mg Cd L¯¹)

Measured concentration (mg Cd L¯¹)

Predicted Cd2⁺ concentration (mg Cd2⁺ L)

0 8 16 32 64 128

0.033 ± 0.01 8.83 ± 0.31 16.77 ± 0.51 30.79 ± 1.03 61.00 ± 1.99 119.37 ± 3.68

0.002 0.58 1.11 2.03 4.03 7.88

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Fig. 1. Cd accumulation in Amphibola crenata viscera, foot muscle and remaining tissues following an acute 48-h waterborne exposure. Plotted values represent mean ± SEM of 6 replicates. Bars sharing lowercase letters are not significantly different with respect to tissue type within an exposure concentration, while bars sharing uppercase letters are not significantly different between different exposure concentrations, within the same tissue, as determined by twoway ANOVA followed by Tukey post-hoc test at p < 0.05.

and complexation, respectively. Waterborne Cd is only considered to be bioavailable in its charged divalent form (Cd2+), where it gains access to the sensitive internal tissues via dedicated divalent metal ion transporters (e.g. DMT-1), or via calcium uptake pathways, by virtue of its mimicry of divalent calcium (Komjarova and Bury, 2014). Thus, the chemistry of seawater acts to reduce Cd toxicity by the formation of Cd chloride species, or through calcium competition for uptake. There is some evidence to support the suggestion that reduced Cd bioavailability is a factor explaining the differences between A. crenata and freshwater pulmonate snails. Pais (2012) reported soft tissue Cd concentrations of ~800 μg g dry wt−1 at an exposure concentration of 600 μg L−1 in the freshwater snail, and although that study examined bioaccumulation after 96 h, these values are still higher (by around 1.4–2-fold) than those in the current study after 48-h exposure to Cd concentrations as high as 32 mg L−1 (Fig. 1). However, it is unlikely that water chemistry alone explains differences in sensitivity between estuarine and freshwater snails. For example, when A. crenata sensitivity was calculated based on predicted Cd2+ concentration, an LC50 value of 3.33 mg L−1 was determined. While this is considerably lower than the value based on dissolved Cd, it is still an order an order of magnitude greater than that determined for freshwater pulmonates. Consequently, it is likely that species-specific differences may also be involved in the high tolerance of A. crenata to waterborne Cd. In aquatic macrofauna the gills are particularly vulnerable to waterborne toxicants. To a significant extent the vulnerability of this tissue stems from its multifunctional roles, as exposure of the gill to a metal toxicant can disrupt ion transport, waste excretion and respiratory function (McGeer et al., 2012). Amphibola lacks a gill, and instead utilises a lung which functions only in respiration, and which is relatively isolated from the external medium. Although such an explanation does not seem to hold for freshwater pulmonates, it is possible that at least a component of the high tolerance of A. crenata may relate to the absence of a gill (Bennington, 1979). In addition, during the Cd exposure, it was noted that many A. crenata individuals produced mucus. This layer may act as a protective barrier, binding Cd, and reducing exposure through subsequent sloughing (Betzer and Pilson, 1974). Moreover, the presence of an operculum in gastropods may also promote tolerance. Under unfavourable conditions the operculum may close, limiting exposure of soft tissues to waterborne toxicants such as Cd (Bennington, 1979). This behavior was unable to be verified in the current study.

Table 2 Relationship between exposure Cd concentration and bioaccumulation of Cd in different tissues of mud snails. Tissue

Coefficient of determination (R2)

Viscera Foot muscle Remaining tissues

0.645 0.822 0.811

Equation for the line of best fit

y = (− 0.69x)/(37.4 + x) y = 2.7x + 7.7 y = (− 0.2x))/(11.6 + x)

p- value

< 0.001 < 0.001 < 0.001

glucose (52% and 75% respectively) were found in snails exposed to 16 and 32 mg Cd L−1 (one-way ANOVA; p < 0.05). The correlation between haemolymph glucose and Cd exposure concentration was strong and significant (R2 = 0.865, p < 0.01; Table 3). Acute Cd exposure resulted in increased haemolymph protein concentrations in mud snails (Fig. 4C). For example, mud snails exposed to 16 and 32 mg Cd L−1 showed 57% and 85% increases in haemolymph protein levels, significantly elevated relative to the control (one-way ANOVA; p < 0.05). Overall, the haemolymph protein concentration correlated significantly with Cd exposure concentration (R2 = 0.631, p < 0.01; Table 3).

4. Discussion 4.1. Acute lethal toxicity and tissue accumulation The 48-h LC50 of adult A. crenata in the current study was 50.4 mg L−1, a value higher than that recorded previously for a gastropod species. For example, Ramakritinan et al. (2012) reported an LC50 value of 39.8 mg L−1 for the marine snail, Cerithedia cingulata, exposed to Cd at a higher temperature (29 versus 15 °C) and salinity (33 versus 20‰) than that used in the current study. Similarly, A. crenata is considerably more tolerant than freshwater pulmonate snails, with a 96-h LC50 of 0.35 mg Cd L−1 having been reported for Lymnaea stagnalis (Pais, 2012). Many factors are known affect the short-term survival of gastropods to Cd exposure, but salinity is clearly a key determinant. Generally, marine species are more tolerant of Cd than freshwater species (Wang et al., 2010). In seawater, high concentrations of ions such as calcium and chloride can reduce Cd bioavailability by processes of competition 278

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Fig. 2. The effect of an acute (48 h) waterborne Cd exposure on A. crenata oxygen consumption (A); ammonia excretion (B); and O:N ratio (C). Plotted values in (A) and (B) represent mean ± SEM of 6 individuals, while plotted values in (C) represent O:N ratio values calculated from group means. Bars sharing letters are not significantly different as determined by one-way ANOVA followed by Tukey post-hoc test at p < 0.05.

of mechanisms that facilitate a storage role. For example, Cd exposure induces expression of the thiol-rich protein metallothionein in gastropod digestive gland and kidney, which is able to chelate Cd in a biologically inert form (Bebianno and Langston, 1998). In the digestive gland, Cd will associate with phosphate granules, which are also considered to detoxify Cd (Nott et al., 1993).

The results of the present study show that the gonad and digestive gland (grouped together in the current study as “viscera”) accumulated the highest tissue burden of Cd in A. crenata exposed to waterborne Cd. Tissue burden in this compartment and the remaining tissues (i.e. kidney, mantle, remaining digestive tissues and heart) showed saturation with increasing Cd exposure concentration (Table 2). The digestive gland and the kidneys of gastropod molluscs are well recognised as tissues of Cd storage (Nott et al., 1993). These tissues display a number

Table 3 Relationship between physiological and biochemical responses, and tissue Cd concentration (for measures made in viscera, foot or remaining tissues) or Cd exposure concentration (physiological and haemolymph measures). Biomarker Oxygen consumption Ammonia excretion Catalase

Lipid peroxidation

Glycogen

Glucose Protein

Tissue

Coefficient of determination (R2)

Equation for the line best fit

p - value

Viscera Foot muscle Remaining tissues Viscera Foot muscle Remaining tissues Viscera Foot muscle Remainig tissues Haemolymph Haemolymph

0.659 0.576 0.584 0.097 0.140 0.071 0.035 0.022 0.338 0.568 0.109 0.865 0.631

y = −0.2x + 5.1 y = 0.06x + 2.7 y = 0.35x + 29.9 y = 0.08x + 28.8 y = 0.008x + 3.4 y = 0.003x + 3.5 y = 0.005x + 1.3 y = 0.003x + 1.7 y = −0.03x + 33.2 y = −0.10x + 14.1 y = −0.03x + 20.8 y = 0.006x + 0.23 y = 0.25x +10.0

< 0.01 < 0.01 < 0.01 0.16 0.11 0.21 0.38 0.48 < 0.01 < 0.01 0.11 < 0.01 < 0.01

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the oxygen consumption rate of A. crenata decreased with increasing Cd exposure concentration. Previous studies have explained declines in oxygen consumption upon metal exposures as being a function of damage to the respiratory epithelium, or a protective mucus production response, both of which would inhibit oxygen uptake (Wu and Chen, 2004). It is also important to note that in bivalve molluscs, shell closure in response to Cd exposure has been offered as an explanation for a decrease in feeding activity (Chandurvelan et al., 2012), and if A. crenata closed their opercula in response to Cd, preventing access of the lung to the oxygenated medium, then this would also decrease oxygen consumption. Changes in oxygen transport entities and oxygen demand could also explain the decline in oxygen consumption with waterborne Cd exposure in A. crenata. For example, in a study on the effects of Cd in the estuarine crab, Chasmagnathus granulata, it was suggested that Cd may alter polymerisation of haemocyanin, thereby reducing haemocyanin oxygen affinity (Rodriguez et al., 2001). Effects on oxygen demand have also been observed, with a significant decline in respiration observed in Cd-exposed mitochondria isolated from a bivalve mollusc, likely due to effects of the metal on mitochondrial function (Sokolova, 2004). All of these factors may have some utility for explaining the observed respiratory impairment in A. crenata in the current study, but identifying the specific mechanism(s) of effect requires further research. Ammonia is the final product of protein catabolism in most aquatic animals (Mayzaud and Conover, 1988). The increase of ammonia excretion in A. crenata exposed to Cd (Fig. 2B), is indicative of an increase in the catabolism of amino acids or other nitrogenous compounds, and is consistent with previous findings in Cd-exposed molluscs (e.g. Chandurvelan et al., 2012). A stimulation in protein catabolism could reflect either enhanced degradation of proteins damaged by Cd exposure (e.g. Tamas et al., 2014), or a change in energy metabolism as a result of Cd exposure. This latter hypothesis is supported by the significant increase in haemolymph protein in Cd-exposed A. crenata (Fig. 4C). This suggests that protein is being mobilised for use as an energy substrate in this species, and is consistent with similar findings in Cd-exposed bivalve molluscs (Chandurvelan et al., 2013; Spann et al., 2011). Further support for the hypothesis that there is a switch to protein metabolism in Cd-exposed mud snails, is provided by changes in the O:N ratio. Hypothetically, O:N ratios less than 16 represent a protein-dominated catabolism, while values higher than 24 represent the predominant use of lipids and carbohydrates as energy substrates (Mayzaud and Conover, 1988). In control animals in the current study, the O:N ratio of 50 was therefore reflective of a reliance on lipids and/ or carbohydrates as fuels. However, the O:N ratio values for A. crenata exposed to Cd were less than 16, thus indicating an increased reliance on protein catabolism. Widdows (1978) suggested that an O:N ratio of 7 or less would be reflective of stress condition. This was the case for A. crenata exposed to waterborne Cd concentrations of 16 and 32 mg L−1, and similar results have been found for juvenile prawns (Exopalaemon carinicauda; Zhang et al., 2014) and green-lipped mussels (Perna canaliculus: Chandurvelan et al., 2012), exposed to Cd.

Fig. 3. The effect of an acute (48 h) waterborne Cd exposure on A. crenata catalase activity (A) and lipid peroxidation (B) in viscera, foot muscle and remaining tissues. Plotted values represent mean ± SEM of 6 replicates. Bars sharing lowercase letters are not significantly different with respect to tissue type within an exposure concentration, while bars sharing uppercase letters are not significantly different between different exposure concentrations, within the same tissue, as determined by two-way ANOVA followed by Tukey post-hoc test at p < 0.05.

4.2. Energetic responses In the current study, biochemical and physiological endpoints were measured in Amphibola individuals that survived a 48-h exposure to Cd. This does not, however, mean that changes in these endpoints are necessarily representative of mechanisms that facilitate tolerance. In order to determine whether changes in biochemical and physiological responses contributed towards survival, the ultimate fate of those individuals would have to be determined. Measured effects could, for example, be changes that precede eventual mortality and thus do not benefit the animal. Nevertheless, the responses measured do reflect the effects of Cd exposure, and thus are likely to be informative of pathways by which this important contaminant exerts toxicity. Further research examining responses at more environmentally realistic exposure concentrations will be required to determine the functional importance of Cd-induced changes in parameters such as bioenergetics. Pulmonate snails are capable of bimodal respiration, and when submerged, perform gas exchange across the cutaneous surface. Amphibola crenata displays equivalent rates of oxygen consumption in air and water (Shumway, 1981), suggesting that the determination of respiratory rates and bioenergetics in an aquatic setting provides a valid assessment of the impact of Cd on these endpoints. In the present study,

4.3. Biochemical endpoints Catalase is an antioxidant defence enzyme responsible for the reduction of H2O2 to water and oxygen (Lushchak, 2011). As H2O2 is the main source of the hydroxyl radical, the most reactive and toxic reactive oxygen species (ROS) in living cells, catalase is considered an important biomarker of oxidative stress, and has been extensively investigated in sentinel species, such as mussels, exposed to Cd (Chandurvelan et al., 2013; Koutsogiannaki et al., 2015). The present study showed that catalase activity was induced in viscera of mud snails exposed to waterborne Cd (Fig. 3A). This suggests that in this tissue Cd perturbed ROS metabolism, resulting in a stimulation of catalase activity. This is an effect that has been seen previously in digestive 280

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Fig. 4. The effect of an acute (48 h) Cd exposure on A. crenata tissue glycogen (A); haemolymph glucose (B); and haemolymph protein (C). Plotted values represent mean ± SEM of 6 replicates. In Panel A, bars sharing lowercase letters are not significantly different with respect to tissue type within an exposure concentration, while bars sharing uppercase letters are not significantly different between different exposure concentrations, within the same tissue, as determined by two-way ANOVA followed by Tukey HSD post-hoc test at p < 0.05. In Panels B and C bars sharing letters are not significantly different as determined by one-way ANOVA, followed by Tukey post-hoc test at p < 0.05.

the present study, visceral and foot muscle glycogen levels dropped significantly in all Cd treatments. This could be attributed to the utilisation of glycogen reserves by the organism through glycogenesis in order to meet energy demand during stress. Depletion of glycogen in aquatic organisms due to waterborne Cd toxicity has been reported in a number of studies (Leung and Furness, 2001; Ansaldo et al., 2006; Chandurvelan et al., 2013). It is, however, worth noting that all experimental animals had been fasted for 96 h by the time of tissue sampling. It is therefore possible that the responses of tissue glycogen observed in the presence of Cd could have been exacerbated by the absence of external nutrient sources. Haemolymph glucose concentrations in aquatic organisms have been recognised as a potential biomarker of a variety of anthropogenic stressors, including trace metals (Hall and van Ham, 1998; Lorenzon, 2005). Increases in glucose haemolymph, such as those observed following Cd exposure in the current study, are likely to be derived from stored glycogen (Cameselle et al., 1980), and this would be consistent with the finding of a reduction in tissue glycogen in A. crenata. However, given that physiological data suggested an increased reliance on protein metabolism in Cd-exposed A. crenata, it would seem likely that at least a component of the increase in haemolymph glucose was through the conversion of glucogenic amino acids to glucose, thereby accounting for the rise in ammonia excretion rate. This would also be consistent with the observed increase in haemolymph protein induced by Cd in A. crenata (Fig. 4C). Lorenzon et al. (2000) found similar

(Chandurvelan et al., 2013) and reproductive (Peng et al., 2015) tissues (i.e. those tissues which constitute the viscera of Amphibola) upon Cd exposure in other marine invertebrates. In contrast, no significant change in catalase activity was found in foot muscle and remaining tissues. This likely reflects the higher Cd accumulation in viscera relative to the other tissues, a finding supported by the positive correlation between Cd burden and catalase activity in this tissue (Table 3). Impairment of the delicate balance that exists between ROS production, and mechanisms that exist to detoxify ROS, can result in oxidative damage, such as lipid peroxidation (Lushchak, 2011). The lack of any significant Cd-induced lipid peroxidation in the current study (Fig. 3B) could be explained a number of ways. One possibility is that Cd did not initiate an imbalance in the relationship between ROS production and ROS scavenging. Alternatively, the enhanced activity of antioxidant defences, such as that seen in the viscera for catalase, were sufficient to prevent damage. Similar findings have been observed in Cd-exposed bivalve molluscs (Viarengo et al., 1999; Chandurvelan et al., 2013). It is also possible that lipid peroxidation was not induced over the relatively short time-frame of the current study (48-h), and that a longer exposure period, or a post-exposure sampling point may have provided a different outcome. Gastropod snails store large quantities of glycogen in their tissues, likely as an energy reserve (Bennington, 1979; Ansaldo et al., 2006). Changes in tissue glycogen can therefore be used as an endpoint that reflects the energy demand of an organism under stress conditions. In 281

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changes in haemolymph glucose to those described here for Cd-exposed mud snails, in the shrimp Palaemon elegans treated with sublethal concentrations of Cd for 24 h. Moreover, Jiang et al. (2013) found hyperglycaemic responses in freshwater crayfish, Cherax quadricarinatus for up to 96 h of Cd exposure. However, hyperglycaemia, has not been reported in all studies. For example, Bislimi et al. (2013) found a hypoglycaemia in the garden snail, Helix pomatia, exposed to industrial pollution. The effects of anthropogenic stressors on haemolymph glucose may, therefore, vary depending on the species.

quantities of protein utilising the principle of protein-dye binding. Anal. Biochem. 72, 248–254. Brix, K.V., Esbaugh, A.J., Grosell, M., 2011. The toxicity and physiological effects of copper on the freshwater pulmonate snail, Lymnaea stagnalis. Comp. Biochem. Physiol. C 154, 261–267. Butler, C.A., Timperley, M.H., 1996. Fertilised farmland as a source of cadmium in oysters. Sci. Total Environ. 181, 31–44. Cameselle, J., Sanchez, J., Carrion, A., 1980. The regulation of glycolysis in the hepatopancreas of the sea mussel Mytilus edulis L. Comp. Biochem. Physiol. B 65, 95–102. Chabicovsky, M., Klepal, W., Dallinger, R., 2004. Mechanisms of cadmium toxicity in terrestrial pulmonates: programmed cell death and metallothionein overload. Environ. Toxicol. Chem. 23, 648–655. Chandran, R., Sivakumar, A.A., Mohandass, S., Aruchami, M., 2005. Effect of cadmium and zinc on antioxidant enzyme activity in the gastropod, Achatina fulica. Comp. Biochem. Physiol. C 140, 422–426. Chandurvelan, R., Marsden, I.D., Gaw, S., Glover, C.N., 2012. Impairment of green-lipped mussel (Perna canaliculus) physiology by waterborne cadmium: relationship to tissue bioaccumulation and effect of exposure duration. Aquat. Toxicol. 124–125, 114–124. Chandurvelan, R., Marsden, I.D., Gaw, S., Glover, C.N., 2013. Biochemical biomarker responses of green-lipped mussel, Perna canaliculus, to acute and subchronic waterborne cadmium toxicity. Aquat. Toxicol. 140–141, 303–313. Chandurvelan, R., Marsden, I.D., Gaw, S., Glover, C.N., 2017. Acute and sub-chronic effects of sub-lethal cadmium exposure on energy metabolism in the freshwater shrimp, Paratya curvirostris. Ecotoxicol. Environ. Saf. 135, 60–67. Chandurvelan, R., Marsden, I.D., Glover, C.N., Gaw, S., 2015. Assessment of a mussel as a metal bioindicator of coastal contamination: relationships between metal bioaccumulation and multiple biomarker responses. Sci. Total Environ. 511, 663–675. Chandurvelan, R., Marsden, I.D., Glover, C.N., Gaw, S., 2016. Biomarker responses of mussels exposed to earthquake disturbances. Estuar. Coast. Shelf Sci. 182, 98–111. Charan-Dixon, H., Gaw, S., Goldstien, S.J., Glover, C.N., 2017. Effects of waterborne cadmium on energy metabolism in the tropical sea cucumber, Stichopus horrens, and a comparison of tissue-specific cadmium accumulation with the temperate sea cucumber Australostichopus mollis. Ecotoxicol. Environ. Saf. 141, 1–8. Craw, D., Brown, K., Webster-Brown, J., 2005. Metals derived from gold mining and geothermal sources. In: Moore, T.A., Black, A., Centeno, J.A., Harding, J.S., Trumm, D.A. (Eds.), Metal contaminants in New Zealand. Resolutionz press, Christchurch NZ, pp. 231–245. Das, S., Khangarot, B.S., 2010. Bioaccumulation and toxic effects of cadmium on feeding and growth of an Indian pond snail Lymnaea luteola L. under laboratory conditions. J. Hazard. Mater. 182, 763–770. Di Toro, D.M., Allen, H.E., Bergman, H.L., Meyer, J.S., Paquin, P.R., Santore, R.C., 2001. Biotic ligand model of the acute toxicity of metals. 1. Technical basis. Environ. Toxicol. Chem. 20, 2383–2396. Eisler, R., 1985. Cadmium Hazards to Fish, Wildlife, and Invertebrates: a Synoptic Review. Contaminant Hazard Reviews, Laurel, MD, pp. 1–35. Grosell, M., Gerdes, R.M., Brix, K.V., 2006. Chronic toxicity of lead to three freshwater invertebrates- Brachionus calyciflorus, Chironomus tentans, and Lymnaea stagnalis. Environ. Toxicol. Chem. 25, 97–104. Gustafsson, J.P., 2012. Visual MINTEQ, 3.0 ed. KTH Royal Institute of Technology, Department of Land and Water Resources Engineering, Stockholm, Sweden. Hall, M.R., van Ham, E.H., 1998. The effects of different types of stress on blood glucose in the giant tiger prawn Penaeus monodon. J. World Aquacult. Soc. 29, 290–299. Hubertz, E.D., Cahoon, L.B., 1999. Short-term variability of water quality parameters in two shallow estuaries of North Carolina. Estuaries 22, 814–823. Jiang, Q.C., Dilixiati, A., Zhang, C., Liu, X.-Z., Huang, W.T., Lv, L.L., Wang, Q., Yang, J.X., 2013. Metabolic and antioxidant responses in juveniles of Cherax quadricarinatus under acute cadmium stress. J. Crustac. Biol. 33, 552–556. Koutsogiannaki, S., Franzellitti, S., Kalogiannis, S., Fabbri, E., Dimitriadis, V.K., Kaloyianni, M., 2015. Effects of cadmium and 17 beta-estradiol on Mytilus galloprovincialis redox status. prooxidant-antioxidant balance (PAB) as a novel approach in biomonitoring of marine environments. Mar. Environ. Res. 103, 80–88. Komjarova, I., Bury, N.R., 2014. Evidence of common cadmium and copper uptake routes in zebrafish Danio rerio. Environ. Sci. Technol. 48, 12946–12951. Korte, F., 1983. Ecotoxicology of cadmium: general overview. Ecotoxicol. Environ. Saf. 7, 3–8. Lagadic, L., Caquet, T., Ramade, F., 1994. The role of biomarkers in environmental assessment (5). Invertebrate populations and communities. Ecotoxicology 3, 193–208. Leung, K.M.Y., Furness, R.W., 2001. Survival, growth, metallothionein and glycogen levels of Nucella lapillus (L.) exposed to sub-chronic cadmium stress: the influence of nutritional state and prey type. Mar. Environ. Res. 2, 173–194. Lorenzon, S., 2005. Hyperglycemic stress response in crustacea. Invert. Surviv. J. 2, 132–141. Lorenzon, S., Francese, M., Ferrero, E.A., 2000. Heavy metal toxicity and differential effects on the hyperglycemic stress response in the shrimp Palaemon elegans. Arch. Environ. Contam. Toxicol. 39, 167–176. Lushchak, V.I., 2011. Environmentally induced oxidative stress in aquatic animals. Aquat. Toxicol. 101, 13–30. Marsden, I.D., Baharuddin, N., 2015. Gastropod growth and survival as bioindicators of stress associated with high nutrients in the intertidal of a shallow temperate estuary. Estuar. Coast. Shelf Sci. 156, 175–185. Marsden, I.D., Swinscoe, I., 2014. Does population structure and growth of an intertidal pulmonate snail reflect environmental conditions within a small estuary? Hydrobiologia 724, 141–155. Matthiessen, P., Law, R.J., 2002. Contaminants and their effects on estuarine and coastal organisms in the United Kingdom in the late twentieth century. Environ. Pollut. 120, 739–757.

4.4. Conclusions In general, estuarine gastropods are tolerant of environmental extremes, including fluctuations in salinity, dissolved oxygen, temperature, osmolality, and the availability of water (Shumway and Marsden, 1982). In the current study, we show that this high tolerance extends to withstanding the lethal impacts of the trace metal Cd, at least over acute (48-h) exposures, in the estuarine pulmonate mud snail Amphibola crenata. This high tolerance contrasts distinctly with the high sensitivity of freshwater pulmonates to Cd (e.g. Das and Khangarot, 2010), even when accounting for water chemistry and the resulting changes in Cd speciation. This suggests that low tolerance to Cd is not solely a function of being a pulmonate snail, but relies on specific physiological characteristics of the different species. This is supported by bioaccumulation data, which show a much greater relative capacity for metal bioaccumulation in the freshwater pulmonates (e.g. Pais, 2012), than in A. crenata. In the current study, even though Cd exposure concentrations were significantly in excess of those likely to be found in natural settings, and Cd increased significantly in snail tissues, there was no increase in oxidative stress over the 48-h exposure, likely due to the stimulated activity of antioxidant pathways such as catalase. There were, however, significant changes in energy metabolism, suggesting that chronic Cd exposures would have a more deleterious effect on mud snail health. We also recognise that in the environment, Cd uptake can occur through many routes including through diet and sediment (McGeer et al., 2012), which could have distinct effects on bioaccumulation patterns and toxicity. Acknowledgments This study was funded by the Brian Mason Scientific and Technical Trust and University of Canterbury (E6573). The authors thank Jonathan Hill, Robert Stainthorpe and Jan McKenzie for their valuable assistance in the laboratory and field. CNG is supported by a Campus Alberta Innovates Program Research Chair. References Ansaldo, M., Nahabedian, D.E., Holmes-Brown, E., Agote, M., Ansay, C.V., Guerrero, N.R., Wider, E.A., 2006. Potential use of glycogen level as biomarker of chemical stress in Biomphalaria glabrata. Toxicology 224, 119–127. Barbieri, E., Paes, E.T., 2011. The use of oxygen consumption and ammonium excretion to evaluate the toxicity of cadmium on Farfantepenaeus paulensis with respect to salinity. Chemosphere 84, 9–16. Bebianno, M.J., Langston, W.J., 1998. Cadmium and metallothionein turnover in different tissues of the gastropod Littorina littorea. Talanta 46, 301–313. Bennington, S.L., 1979. Some aspects of the biology and distribution of Amphibola crenata (Gastropoda: pulmonata) (Unpublished Ph.D. Thesis). University of Canterbury, Christchurch, New Zealand. Besser, J.M., Dorman, R.A., Hardesty, D.L., Ingersoll, C.G., 2016. Survival and growth of freshwater pulmonate and nonpulmonate snails in 28-day exposures to copper, ammonia, and pentachlorophenol. Arch. Environ. Contam. Toxicol. 70, 321–331. Betzer, S.B., Pilson, M.E.Q., 1974. The seasonal cycle of copper concentration in Busycon canaliculatum L. Biol. Bull. 146, 165–175. Bislimi, K., Behluli, A., Halili, J., Mazreku, I., Osman, F., Halili, F., 2013. Comparative analysis of some biochemical parameters in haemolymph of garden snail (Helix pomatia L.) of the Kostriot and Ferizaj Regions, Kosovo. Int. J. Eng. Appl. Sci. 4, 11–18. Böning, P., Brumsack, H.J., Bottcher, M.E., Schnetger, B., Kriete, C., Kallmeyer, J., Borchers, S.L., 2004. Geochemistry of Peruvian near-surface sediments. Geochim. Cosmochim. Acta 68, 4429–4451. Bradford, M.M., 1976. A rapid and sensitive method for the quantification of microgram

282

Ecotoxicology and Environmental Safety 158 (2018) 274–283

N.A.L. De Silva et al.

Shuhaimi-Othman, M., Nadzifah, Y., Nur-Amalina, R., Umirah, N.S., 2013. Deriving freshwater quality criteria for copper, cadmium, aluminum and manganese for protection of aquatic life in Malaysia. Chemosphere 90, 2631–2636. Shuhaimi-Othman, M., Nur-Amalina, R., Nadzifah, Y., 2012. Toxicity of metals to a freshwater snail, Melanoides tuberculata. Sci. World J. 2012, 125785. Shumway, S.E., 1981. Factors affecting oxygen consumption in the marine pulmonate Amphibola crenata (Gmelin, 1791). Biol. Bull. 160, 332–347. Shumway, S.E., Marsden, I.D., 1982. The combined effects of temperature, salinity, and declining oxygen tension on oxygen consumption in the marine pulmonate Amphibola crenata (Gmelin, 1791). J. Exp. Mar. Biol. Ecol. 61, 133–146. Sokolova, I.M., 2004. Cadmium effects on mitochondrial function are enhanced by elevated temperatures in a marine poikilotherm, Crassostrea virginica Gmelin (Bivalvia: ostreidae). J. Exp. Biol. 207, 2639–2648. Spann, N., Aldridge, D.C., Griffin, J.L., Jones, O.A.H., 2011. Size-dependent effects of low level cadmium and zinc exposure on the metabolome of the Asian clam, Corbicula fluminea. Aquat. Toxicol. 105, 589–599. Tamas, M.J., Sharma, S.K., Ibstedt, S., Jacobson, T., Christen, P., 2014. Heavy metals and metalloids as a cause for protein misfolding and aggregation. Biomolecules 4, 252–267. Viarengo, A., Burlando, B., Cavaletto, M., Marchi, B., Ponzano, E., Blasco, J., 1999. Role of metallothionein against oxidative stress in the mussel Mytilus galloprovincialis. Am. J. Physiol. 277, R1612–R1619. Wang, Z., Yan, C., Kong, H., Wu, D., 2010. Mechanisms of Cadmium Toxicity to Various Trophic Saltwater Organisms. Nova Science Publishers, Inc, New York. Widdows, J., 1978. Physiological indices of stress in Mytilus edulis. J. Mar. Biol. Assoc. U. K. 58, 125–142. Wu, J.P., Chen, H.C., 2004. Effects of cadmium and zinc on oxygen consumption, ammonium excretion, and osmoregulation of white shrimp (Litopenaeus vannamei). Chemosphere 57, 1591–1598. Zhang, C., Li, F., Xiang, J., 2014. Acute effects of cadmium and copper on survival, oxygen consumption, ammonia-N excretion, and metal accumulation in juvenile Exopalaemon carinicauda. Ecotoxicol. Environ. Saf. 104, 209–214.

Mayzaud, R., Conover, R.J., 1988. O:N atomic ratio as a tool to describe zooplankton metabolism. Mar. Ecol. Prog. Ser. 45, 289–302. McDowell, R.W., 2010. Is cadmium loss in surface runoff significant for soil and surface water quality: a study of flood-irrigated pastures? Wat. Air Soil Pollut. 209, 133–142. McGeer, J.C., Niyogi, S., Smith, D.S., 2012. Cadmium. In: Wood, C.M., Farrell, A.P., Brauner, C.J. (Eds.), Homeostasis and Toxicology of Non-essential Metals, Fish Physiology 31B. Elsevier, London, UK, pp. 125–184. McRae, N.K., Gaw, S., Glover, C.N., 2016. Mechanisms of zinc toxicity in the galaxiid fish, Galaxias maculatus. Comp. Biochem. Physiol. C 179, 184–190. McRae, N.K., Gaw, S., Glover, C.N., 2018. Effects of waterborne cadmium on metabolic rate, oxidative stress, and ion regulation in the freshwater fish, inanga (Galaxias maculatus). Aquat. Toxicol. 194, 1–9. Nair, A.R., Degheselle, O., Smeets, K., Van Kerkhove, E., Cuypers, A., 2013. Cadmiuminduced pathologies: where is the oxidative balance lost (or not)? Int. J. Mol. Sci. 14, 6116–6143. Nott, J.A., Bebianno, M.J., Langston, W.J., Ryan, K.P., 1993. Cadmium in the gastropod Littorina littorea. J. Mar. Biol. Assoc. U.K. 73, 655–665. Pais, N.M., 2012. Studies on waterborne cadmium exposure to Lymnaea stagnalis in varying water qualities and the development of a novel tissue residue approach (Unpublished M.Sc. Thesis). Wilfrid Laurier University, Waterloo, ON, Canada. Peng, C., Zhao, X.G., Han, Y., Shi, W., Liu, S.X., Liu, G.X., 2015. Toxic effects of chronic sub-lethal Cu2+, Pb2+ and Cd2+ on antioxidant enzyme activities in various tissues of the blood cockle, Anadara granosa. J. Residuals Sci. Technol. 12, 125–131. Ramakritinan, C., Chandurvelan, R., Kumaraguru, A.K., 2012. Acute toxicity of metals: cu, Pb, Cd, Hg and Zn on marine molluscs, Cerithedia cingulata G., and Modiolus philippinarum H. Ind. J. Geo-Mar. Sci. 41, 141–145. Ridgway, J., Shimmield, G., 2002. Estuaries as repositories of historical contamination and their impact on shelf seas. Estuar. Coast. Shelf Sci. 55, 903–928. Rodriguez, E.M., Bigi, R., Medesani, D.A., Stella, V.S., Greco, L.S., Moreno, P.A., Monserrat, J.M., Pellerano, G.N., Ansaldo, M., 2001. Acute and chronic effects of cadmium on blood homeostasis of an estuarine crab, Chasmagnathus granulata, and the modifying effect of salinity. Braz. J. Med. Biol. Res. 34, 509–518.

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